Herbicides Environmental Impact Studies and Management Approaches Part 4 pptx

Use of Sugar Beet as a Bioindicator Plant
for Detection of Flucarbazone and Sulfentrazone Herbicides in Soil

49
type and are generally lower in sandy soils of low organic matter content and high pH
(Jourdan et al. 1998; Eliason et al. 2004; Szmigielski et al. 2009).
Since flucarbazone and sulfentrazone decrease both root and shoot length of sensitive plants
such as sugar beet, sequential or simultaneous applications of these two herbicides could
potentially result in herbicide interactions.
3. Flucarbazone and sulfentrazone interactions
Repeated applications of herbicides with the same mode of action have resulted in weeds
developing resistance (Vencill et al. 2011; Colborn & Short 1999; Whitcomb 1999). Using
herbicides with different mode of action either applied as pre-mixed combinations or
applied in rotation reduces problems related to weed resistance and consequently improves
weed control. However, combinations of herbicides are generally chosen to improve the
spectrum of weed control without prior knowledge of the possible consequences of the
interactions between herbicides (Zhang et al. 1995). The outcome of the interactions may be
synergistic, antagonistic or additive depending on whether the combined effect on the target
plants is greater, less than, or equal to the summed effect of the herbicides applied alone
(Colby 1967; Nash 1981). A synergistic interaction occurs when the activity of two herbicides
is more phytotoxic than either herbicide applied singly. A synergistic effect is beneficial in
that it provides more effective weed control at lower herbicide concentrations; however it
may also cause injury to sensitive rotational crops if the synergism of the two residual
herbicides is not known (Zhang et al. 1995). In an additive interaction, also called “herbicide
stacking” (Johnson et al. 2005), the injury observed in the target plants is the sum activity of
the combined herbicides. With an antagonistic interaction, the efficacy of the combined
herbicides is reduced and consequently results in decreased weed control but can also help
to avoid unwanted crop injury (Zhang et al. 1995).
To examine interactions between soil-incorporated flucarbazone and sulfentrazone, we
evaluated the combined effect of these two herbicides on sugar beet root and shoot
inhibition. Root length inhibition was assessed in soil that was spiked with mixtures

consisting of flucarbazone in the range from 0 to 15 ppb with sulfentrazone added at 50 ppb
level, while shoot length inhibition was evaluated in soil that was amended with mixtures
consisting of sulfentrazone in the range from 0 to 200 ppb with flucarbazone added at 6 ppb
level. The expected inhibition was calculated using Colby’s formula (Colby 1967):
E = X + Y – XY/100 (3)
where X is the plant growth inhibition (%) due to compound A and Y is the plant growth
inhibition (%) due to compound B; comparing expected inhibition to the observed inhibition
allows the nature of interactions to be revealed. The combined effect of flucarbazone and
sulfentrazone was additive: the observed and expected root length inhibition of sugar beet
in response to flucarbazone in combination with sulfentrazone were similar (Fig. 4a), as
were the observed and the expected shoot length inhibition due to sulfentrazone in
combination with flucarbazone (Fig. 4b). I
50
values for observed and expected responses
were not different at 0.05 level based on the asymptotic z-test. The additive effect of
flucarbazone and sulfentrazone will help in weed control but may also increase risk of
injury to rotational crops that are sensitive to both these herbicides.

Herbicides – Environmental Impact Studies and Management Approaches

50
Flucarbazone (ppb) + sulfentrazone (50 ppb)
0246810121416
Root length inhibition (%)
0
20
40
60
80
100

Observed root length inhibition
Expected root length inhibition

Sulfentrazone
(pp
b
)
+ flucarbazone
(
6
pp
b
)
050100150200
Shoot length inhibition (%)
0
20
40
60
80
100
Observed shoot length inhibition
Expected shoot length inhibition

Fig. 4. (a) Root length inhibition of sugar beet in response to increasing concentration of
flucarbazone in combination with 50 ppb sulfentrazone, and (b) shoot length inhibition of
sugar beet in response to increasing concentration of sulfentrazone in combination with 6
ppb flucarbazone.
4. Effect of ammonium containing fertilizer on sugar beet bioassay
Typically plant response that is measured in a bioassay is not specific to one source. The lack

of specificity may be desirable in that the presence of residues of all herbicides that
detrimentally affect the same plant parameter are detected. However, other soil applied
chemicals apart from herbicides may also alter the parameter measured in a bioassay and
may change the outcome of the bioassay. We have reported that the detection of ALS-
inhibiting herbicides in soil using a mustard root bioassay is influenced by N-fertilizer as
mustard root length is shortened in response to ammonium ions (Szmigielski et al. 2011).
Ammonium toxicity to plants is common and a change in root/shoot ratio is one of the
symptoms of NH
4
+
toxicity (Britto & Kronzucker 2002).
To assess the effect of N-fertilizer on sugar beet roots and shoots, and consequently on
flucarbazone and sulfentrazone detection in soil, ammonium nitrate was added to soil in the
range from 0 to 200 ppm N, and root and shoot length was measured. Ammonium nitrate
significantly reduced root length of sugar beet but the shoot length inhibition due to
ammonium nitrate was very small and was less than 20% at the highest ammonium nitrate
concentration tested (Fig. 5).
The combined response of sugar beet roots to flucarbazone and ammonium nitrate was
examined by growing sugar beet plants in soil that was spiked with flucarbazone in the
range of 0 to 15 ppb and mixed with ammonium nitrate added at 50 ppm N. The expected
response due to flucarbazone in combination with ammonium nitrate was calculated using
equation [3]. Since the expected root length inhibition was the same as the observed (Fig. 6),
the combined effect of flucarbazone and N-fertilizer on sugar beet root length is additive.
Thus, root length reduction of sugar beet that is measured in a soil that received a recent
application of ammonium containing or ammonium producing fertilizer may be
(a)
(b)
Use of Sugar Beet as a Bioindicator Plant
for Detection of Flucarbazone and Sulfentrazone Herbicides in Soil

51
misinterpreted as reduction due to herbicide residues and may yield false positive results.
Because N-fertilizer interferes with the sugar beet root length bioassay, preferably soil
sampling for the detection of residual herbicides should be completed preplant and before
N-fertilizer field application, or at the end of the growing season.

-10
0
10
20
30
40
50
60
70
6.25 12.5 25 50 100 200
N concentration (ppm)
Inhibition (%)
Shoot length inhibition
Root length inhibition

Fig. 5. Effect of increasing ammonium nitrate concentration in soil on shoot and root
inhibition of sugar beet plants.

Flucarbazone (ppb) + NH
4
NO
3
(50 ppm N)
0 2 4 6 8 10 12 14 16

Root length inhibition (%)
0
20
40
60
80
100
Observed root length inhibition
Expected root length inhibition

Fig. 6. Root length inhibition of sugar beet in response to increasing concentration of
flucarbazone in combination with 50 ppm N added as ammonium nitrate.

Herbicides – Environmental Impact Studies and Management Approaches

52
5. Effect of landscape position on phytotoxicity and dissipation of
flucarbazone and sulfentrazone
Farm fields with irregular rolling topography of low hills and shallow depressions are
typical on the Canadian prairies (Fig. 7). Low-slope soils from depressions in the landscape
typically have higher organic matter and clay contents and lower pH than up-slope soils
from elevated parts of the terrain (Schoenau et al. 2005; Moyer et al. 2010). Furthermore,
low-slope areas in the field generally have higher moisture content as a result of water
accumulating in the depressions, while the up-slope areas are drier due to water runoff.

Fig. 7. Undulating landscape comprised of knolls and depressions in southwestern
Saskatchewan (source: Geological Survey Canada).
Phytotoxicity of ALS- and protox-inhibiting herbicides is soil dependent, and the effect of
organic matter, clay and soil pH on adsorption and bioavailability of these herbicides is well
documented (Thirunarayanan et al. 1985; Renner et al. 1988; Che et al. 1992; Wang & Liu

1999; Wehtje et al. 1987; Grey et al. 1997; Szmigielski et al. 2009). Typically organic matter
and clay decrease the concentration of bioavailable herbicide through adsorption of
herbicide molecules to the reactive functional groups and colloidal surfaces. At alkaline soil
pH, adsorption of weak acidic herbicides tends to decrease due to increased herbicide
solubility in soil solution and due to repulsion of anionic herbicide molecules from
negatively charged soil particles.
Dissipation of ALS- and protox-inhibiting herbicides in soil is governed by microbial and
chemical processes. Microbial degradation is the primary mechanism as dissipation has
been shown to be faster in non-sterile soil than in autoclaved soil (Joshi et al. 1985; Ohmes at
al. 2000; Brown 1990). The dissipation rate of ALS- and protox-inhibiting herbicides varies
with soil type and environmental conditions. Generally high organic matter content, high
clay content and low soil pH decrease the dissipation rate by reducing the amount of
herbicide available in soil solution for decomposition (Eliason et al. 2004; Goetz et al. 1990;
Beckie & McKercher 1989; Ohmes et al. 2000; Grey et al. 2007; Main et al. 2004). Microbial
and chemical decomposition both depend on soil water and temperature with faster
dissipation occurring in moist and warm soils (Beckie & McKercher 1989; Joshi et al. 1985;
Use of Sugar Beet as a Bioindicator Plant
for Detection of Flucarbazone and Sulfentrazone Herbicides in Soil

53
Walker & Brown 1983; Brown, 1990; Thirunarayanan et al. 1985). In flooded (saturated) soils
decomposition may be reduced due to anaerobic conditions.
To examine the effect of landscape position on phytotoxicity and dissipation of flucarbazone
and sulfentrazone, we used two soils that were collected from a farm field with varying
topography in southern Saskatchewan, Canada. Soil from an up-slope position contained
0.9% organic carbon, 31% clay and had pH 7.9, while soil from a low-slope position
contained 1.6% organic carbon, 51% clay and had pH 7.2. Flucarbazone phytotoxicity was
assessed in the range from 0 to 15 ppb by measuring root length inhibition while
sulfentrazone phytotoxicity was determined in the range from 0 to 200 ppb by measuring
shoot length inhibition of sugar beet. Phytotoxicity of flucarbazone (Figure 8a) and of

sulfentrazone (Figure 8b) was higher in the up-slope soil than in the low-slope soil. The I
50

values determined from the dose-response curves were 3.5 and 5.7 ppb for flucarbazone,
and 34.3 and 56.5 ppb for sulfentrazone in the up-slope and low-slope soil, respectively, and
were different at 0.05 level of significance. Thus landscape position in a field has a
considerable effect on bioavailability of flucarbazone and sulfentrazone, and different
herbicide application rates may be required in fields of variable topography to achieve
uniform weed control.
Flucarbazone
(pp
b
)
0246810121416
R
oo
t

l
eng
th

i
n
hibiti
on
(%)
0
20
40

60
80
100
Up-slope soil
Low-slope soil

Sulfentrazone
(pp
b
)
0 50 100 150 200
Shoot length inhibition (%)
0
20
40
60
80
100
Up-slope soil
Low-slope soil

Fig. 8. Dose-response curves for (a) flucarbazone determined by root length, and (b)
sulfentrazone determined by shoot length of sugar beet in soil from two landscape
positions.
Flucarbazone and sulfentrazone dissipation in the two soils was examined under laboratory
conditions of 25 C and moisture content of 85% field capacity. Soils were spiked with 15 ppb
of flucarbazone and separately with 200 ppb of sulfentrazone, and at each sampling time the
residual flucarbazone and sulfentrazone was determined using the sugar beet bioassay.
Flucarbazone and sulfentrazone dissipation followed the bi-exponential decay model
described in detail by Hill & Schaalje (1985):

C = a e
-bt
+ c e
-dt
(4)
where C is herbicide concentration remaining in soil after time t. In the bi-exponential
decay model the dissipation rate is not constant and is fast initially and slow afterward,
(a)
(b)

Herbicides – Environmental Impact Studies and Management Approaches

54
while in the first order decay model (when b = d in equation [4]) the dissipation rate does
not change with time. Flucarbazone and sulfentrazone dissipation was more rapid in the
up-slope soil than in the low-slope soil (Fig. 9a and 9b); flucarbazone half-life was 5 and 8
days, and sulfentrazone half-life was 21 and 90 days in the up-slope and the low-slope
soil, respectively. Thus landscape positions in the field influence persistence of
flucarbazone and sulfentrazone, and consequently may affect the potential for herbicide
carry-over to the next growing season. However, because damage to sensitive rotational
crops occurs when a herbicide is available to plants at harmful concentrations one year
after application, risk of carry-over injury is controlled by the combined effect of herbicide
dissipation and herbicide phytotoxicity, both of which are soil dependent; also the
rotational crop must be susceptible to the residual herbicide concentration at the time of
planting (Hartzler et al. 1989). Although flucarbazone and sulfentrazone persist longer in
soil from depressions in the field, herbicide bioavailability is reduced in this soil, and thus
residual flucarbazone or sulfentrazone may not pose a risk of injury to sensitive crops in
low-slope areas. Predicting carry-over injury due to flucarbazone and sulfentrazone in
farm fields with varying topography is a complex task and all factors that affect herbicide
persistence and bioavailability have to be considered before choosing a rotational crop to

grow.
Days
0 1020304050
Flucarbazone remaining (%)
0
20
40
60
80
100
Up-slope soil
Low-slope soil

Da
y
s
0 50 100 150 200 250 3
0
S
u
lf
en
t
razone rema
i
n
i
ng
(%)
0

20
40
60
80
100
Up-slope soil
Low-slope soil


Fig. 9. Dissipation under laboratory conditions of (a) flucarbazone determined by root
length, and (b) sulfentrazone determined by shoot length of sugar beet in soil from two
landscape positions.
6. Practical considerations
Because sugar beet plants respond both to flucarbazone and sulfentrazone, a sugar beet
bioassay allows for detection of these two herbicides in soil by evaluating both root and
shoot inhibition. Growing sugar beet plants in Whirl-Pak
TM
bags is simple and provides a
convenient method for assessing shoot and root length. Shoots are measured above the soil
level and do not need to be harvested; this helps particularly with measuring shoots that are
short and brittle at phytotoxic sulfentrazone concentrations. Roots are recovered from soil
with water and consequently roots do not get broken or damaged before being measured.
(a) (b)
Use of Sugar Beet as a Bioindicator Plant
for Detection of Flucarbazone and Sulfentrazone Herbicides in Soil

55
Furthermore, as the bioassay is completed before roots grow to the bottom of the bag, root
development in Whirl-Pak
TM

bags is not obstructed.
7. Conclusions
Using the sugar beet bioassay we determined: (1) that while flucarbazone primarily inhibits
root length it also causes shoot reduction and while sulfentrazone primarily inhibits shoot
length it also affects root development, (2) that the combined effect of soil-incorporated
flucarbazone and sulfentrazone on root and shoot length inhibition of sugar beet is additive,
(3) that N-fertilizer reduces root length of sugar beet but has little effect on shoot length and
therefore the presence of freshly applied N-fertilizer may yield false positive results for
flucarbazone residues, and (4) that flucarbazone and sulfentrazone phytotoxicity is higher
and dissipation rate is faster in soils from up-slope than low-slope landscape positions
under identical moisture and temperature conditions.
8. Acknowledgements
The financial support of FMC Corporation Canada, Arysta LifeScience Canada, and NSERC
is gratefully acknowledged.
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4
Investigation of Degradation of Pesticide
Lontrel in Aqueous Solutions
E.A. Saratovskih
Institute of Problem of Chemical Physics,
Russian Academia of Science
Russia
1. Introduction
Development of new efficient methods for purifying water from industrial pollutants
resistant to biodegradation is a challenge because of the current shortage of freshwater
reserves in the world (Legrini et al., 1993; Skurlatov et al., 1994). It is known that application
of pesticides (Ozelenenie, 1984; Bykorez, 1985; Burgelya & Myrlyan, 1985; Patel et al., 1991;
Wan et al., 1994; Arantegui et al., 1995; Fliedner, 1997; Arkhipova et al., 1997), in particular,
Lindane (Fliedner, 1997) and 2,4-D (Arkhipova et al., 1997), results in their long-term
accumulation in soils and water reservoirs. Furthermore, pesticides' residual presence in
foodstuffs and industrial crops causes serious diseases, including hereditary disorders
(Eikhler, 1993; Calvert et al., 2004; Whyatt et al., 2004).
Numerous studies point to long-term preservation of many pesticides in natural waters:
ground, river, or sea (Skurlatov et al., 1994; Yudanova, 1989; Yablokov, 1990). Organic
chemicals, including the insecticide Lindane, were found in the fat of gray whales caught in
the Arctic Ocean (Poliakova et al., 2005).
Lontrel (another commercial name is Clopyralid) is considered to be the herbicide of a wide

range of action and, first, for the defense against weeds in grain crops and can be used in a
mixture with 2M-4X or 2,4-D; LD
50
≈ 5000mg/kg (Mel`nikov, 1987). Literature data on the
mechanism of action of Lontrel are rather insignificant. There are indications (Hall et al.,
1985) that Lontrel exhibits auxine-like activity. The most detailed studies are presented in
our works. Metal complexes of Lontrel were not studied at all, except of our research group.
Formerly (Aliev et al., 1988; Saratovskikh et al., 1989b) we showed that 3,6-Dichloropicolinic
acid (DCPA), the active principle of the herbicide Lontrel, readily formed complexes with
metals, major environmental pollutants, and these complexes were stable under natural
conditions. They are capable of participating in further complex formation with bioactive
ligands due to the filling of the coordination sphere of the metal.
We have proved for the first time that pesticides themselves and their metal complexes
interact with mono-, di-, and polynucleotides (Saratovskikh et al., 1988; 1989a). In all cases,
two- or three-component complex systems are formed. It was shown that the pesticide
complex with adenosine triphosphoric acid is formed due to the protonation of the N-7

Herbicides – Environmental Impact Studies and Management Approaches

60
nitrogen atom of the adenine heterocycle, and the nitrogen atom of the terminal NH
2
group
can simultaneously be bound to the pesticide molecule due to the formation of a hydrogen
bond. The formation of the pesticide complexes with ATP results in an energy deficient in
the tissues of organisms. The effect of pesticides and their metal complexes induces the
energy deficient of the cell, namely, inhibition of energy metabolism due to the formation of
a complex with adenosine triphosphoric acid.

N

N
H
N
N
NHH
HO
O
C
N
Cl
Cl
+
Structure of the [ATF-DCPA] complex. Structure of complex of ε-ATF with metal
complex of DCPA - [ε-ATF-CuL
2
].
We pioneered to show that Lontrel and its metal complexes simultaneously inhibit the
activity of several enzymatic systems in organisms. These are oxidizing enzymatic systems,
for example, NADH-oxidoreductase. In this case, the enzymatic activity is inhibited, first,
due to the complex formation with dinucleotide (coenzyme) NADH. Second, this occurs due
to the formation of the non-productive complex [enzyme-DCPA] in the active center of the
enzyme. For DCPA comparison of inhibition of NADH-oxidoreductase К
i
·= 10
4
М, and
competitive type of inhibition take place. The antireductase activity increases in the order:
metal ion < DCPA < metal complex (Saratovskikh et al., 2005; 2007c).
We shown than DCPA and its metal complexes are bound into complex compounds with
DNA and RNA (Saratovskikh et al., 1989a) and nativity of the DNA double helix is violated

because of complex formation. The direct mutagenic effect was shown by us for the ТА98
Salmonella typhimurium strain, mutations of the reading frame shift type are induced and
promutagenity was revealed (Saratovskikh et al., 2007b).
Therefore, it is important to investigate its exposure to a microbial community of activated
sludge (AS) and to sunlight, i.e., conditions mimicking those occurring in natural surface
water bodies.
The possibility of using UV radiation for the decomposition of various chemical compounds
was shown by several authors (Legrini et al., 1993; Guittonneau et al., 1988; Sundstrom et al.,
1989; Castrantas & Gibilisco, 1990). Ultraviolet purification of water is superseding
conservative chlorine treatment (Skurlatov et al., 1994; Skurlatov & Shtamm, 1997a, 2002).
At present, over 1000 ultraviolet-purification devices of different capacities are in operation
in 35 countries (Skurlatov et al., 1996; Kruithof et al., 1992). They finely replace the old
method of treatment of drinking and waste waters based on the treatment with chlorine. A
comparative estimation of the American specialists (Purus Inc., 1992) of the cost of UV
quanta, ozone, and reagents used in processes of water preparation and water treatment is
as follows: Cl
2
= 0,16$ per 1 mol; О
3
= 0,1$; photons 185, 254 nm (low-pressure Hg lamps,
yield 40%) = 0,025$. The bactericide light quantum turned out to be the cheapest reagent
(Skurlatov & Shtamm, 1997b). Sometimes the UV treatment is combined with the addition of

Investigation of Degradation of Pesticide Lontrel in Aqueous Solutions

61
oxygen or hydrogen peroxide. As shown (Shinkarenko & Aleskovskii, 1982), ozone formed
upon the photochemical oxidation of oxygen dissociates, in turn, to the electron-excited
oxygen atom and oxygen molecule in the singlet state. Having a high oxidizability, they
attack a molecule of the contaminant and oxidize it (partially or completely). The UV

treatment of waste waters produces no dibenzofurans, which are precursors of dioxins as it
takes place for the treatment with ozone only (Beltran et al., 1993). The ultraviolet treatment
of waste waters is well compatible with methods of biological purification from
contaminants (Golubovskaya, 1978). Therefore, study of the products of UV-mediated
DCPA degradation is of practical importance.
The objective of the present study was to investigate the possibility of DCPA degradation by
biochemical and photochemical methods (taken alone or in combination with each other), in
order to develop practical methods of degrading pesticides, the most challenging industrial
pollutants.
Analysis of the composition of the reaction mixture and final products is a complicated task,
which demands up-to-date highly sensitive methods, but it is of environmental importance.
This study concerns the kinetics of DCPA degradation under natural conditions (deep-well,
river, or sea water) and analysis of its products.
2. Materials and methods
2.1 Used substances, concentrations and replicates
3,6-Dichloropicolinic acid (DCPA) was studied, which is the main active principle of Lontrel
and a copper–DCPA complex containing two DCPA molecules per copper atom - CuL
2
.
Experiments were performed with 1.22·10
–3
M DCPA solutions in distilled water, artificially
made sea water, and river water from the Klyazma River. Solutions of CuL
2
with the
concentration 0.4·10
–3
M. The artificial seawater was made from sea salt, containing the
following anions (g/l): Na
+

(10.76), K
+
(0.39), Ca
2+
(0.41), Mg
2+
(1.30), SO
4
2-
(2.70), Cl
–
(19.35),
Br
–
(0.06), and CO
3
2-
(0.07). The concentration of sea salt in the working solution was 35.5 g/l
(Artobolevskii, 1977). River water was sampled near Biokombinat Village, Moscow Region,
in the middle Klyazma River at the depth 1.5–2 m.

N
Cl
Cl
COO
H


N
C

O
Met
O
O
N
C
O
C
L
CL
CL
CL
3,6-Dichloropicolinic acid – DCPA
bis-(3,6-Dichloropicolinato)cuper(II)
CuL
2

All experiments were carried out three times. The measurement errors did not exceed
5-10%.
2.2 The biochemical oxidation of DCPA
Biochemical oxidation of DCPA was studied in 1 l laboratory aeration tanks under
conditions of constant aeration and natural sunlight. Concentrated AS was sampled from an

Herbicides – Environmental Impact Studies and Management Approaches

62
aeration tank of the Treatment plants in Chernogolovka, Moscow oblast. Five samples were
studied, each containing 2 ml of AS. The samples were diluted in laboratory tanks to 1 l with
a peptone medium. Sample 1 was taken as a control, not exposed to the mutagen (N-methyl-
N-nitrosourea, nitrosomethylurea, NMU). NMU was added to samples 2, 3, and 4 to a

concentration of 0.07% (Rapoport, 2010); samples 2 and 4 were exposed to NMU for 6 h,
while sample 3 was exposed to NMU for 18 h. Samples 2–4 were treated with NMU once
more for the same amount of time as in the initial treatment. Sample 4 was then treated the
again after 28 days of the observation, while samples 2 and 3 were treated again after 44
days. Samples 1–3 were supplemented with DCPA at a starting concentration of 1.22·10
–3
M
(0.23 g/l). Sample 4 was supplemented with a Cu(L)
2
. This complex was synthesized
according to (Aliev et al., 1988). We showed previously that this complex was a stronger
herbicide than DCPA itself (Saratovskikh et al., 1990). It forms stable compounds with DNA
(Saratovskikh et al., 1989a), mononucleotides, and dinucleotides (Saratovskikh et al., 1988).
Such complexes can readily form in runoff from farmlands and in some industrial
enterprises’ wastewaters containing Lontrel and copper compounds. Sample 5 was
supplemented with wastewater of a pilot Lontrel mixture, which contained approximately
~10
–3
M of the herbicide. Samples for the tests were taken four or five times per day for the
first 3 days, once a day for the next 10 days, and then once every few days. The observation
was performed for one year. The range of microbial species in the AS was determined
according to (Belyaeva & Gyupter, 1969; Liperovskaya, 1977).
2.3 The photochemical oxidation of DCPA
Photochemical oxidation of DCPA under UV irradiation was performed in a quartz reaction
vessel with internal diameter 2.2 cm and volume 25 ml at 4 cm from the ray emitter (in various
sets of experiments, irradiation was performed with DRSh-1000, DRB-8, or BRA-15 lamps;
250–600 nm; BRA-15 lamp 1.3 mW/cm
2
; Institute of Problems of Chemical Physics, Russia).
The starting DCPA concentration was 5·10

–4
M. Oxygen, ozone, air, or argon were sparged
(bubbled through the solution) with a capillary tube. Samples for measuring concentrations
were taken at 10-min intervals. DCPA concentrations were measured with a Specord UV-VIS
spectrometer (Karl Zeiss, Jena, Germany) according to light absorption at 283 nm.
2.4 Toxicological testings
The samples were subjected to biotests with two sentinel species: the infusorium
Tetrahymena pyriformis and the luminescent bacterium Beneckea harveyi.
2.4.1 The biotest on Tetrahymena pyriformis
Tetrahymena pyriformis were cultivated in a peptone or carbohydrate–salt–yeast extract
media (Yoshioka et al., 1985). Five milliliters of each sample were placed into three test tubes
and one or two drops of a 3- to 4-day-old Tetrahymena culture were inoculated. Settled
aquarium water was used as a control. The division period for T. pyriformis ranges within 4–
6 h. The test lasted for 15 min. A change in the behavior or morphology of infusorian cells in
the samples or their death was indicative of a strong toxic effect. Fifteen minutes after the
mixing, three samples (one drop per sample) were taken from each test tube and fixed with
iodine. Infusorian cells were counted under a microscope. In this method, toxicity
coefficients K were calculated as follows:

Investigation of Degradation of Pesticide Lontrel in Aqueous Solutions

63
K = [(Ac–Aex)/Ac] 100%, (1)
where Ac is the number of cells counted by microscopic examination in the control sample
and Aex is the number of cells in an experimental sample. A sample considered toxic at K >
50%. The statistical significance of values was determined by the Student’s -test.
2.4.2 Biotest on Beneckea harveyi
In a chronic test, the toxicity of a sample was judged from the suppression of cell
propagation after 24 h of incubation. A reduction in cell propagation by 50% or more was
indicative of toxicity. A lyophilized preparation of B. harveyi was stored in a freezing

chamber. Immediately before the experiments, the bacteria were suspended in 3% NaCl. For
testing toxicity, 0.3–0.5 ml of the suspension was added to 0.5 ml of a water sample. The
control experiment was performed with a 0.85% NaCl solution. Measurements were
performed with a BLM-8801 luminometer (SKTB Nauka, USSR) with voltmeter detection.
Toxicity was estimated from the decrease in bioluminescence of a sample relative to the
control. The sample was considered toxic if its bioluminescence decreased by 50% or more.
The level of bacterial luminescence is determined by the intensity of intracellular
metabolism involving the enzyme luciferase. A decrease in luminescence may be related to
either inhibition of the enzyme itself or to an effect of toxic substances on other links in the
metabolic chain. The toxicity coefficient was calculated as:
T= [(Ic–Iex)/Ic] 100%, (2)
where Ic is the bioluminescence intensity in the control and Iex is the bioluminescence
intensity in the sample tested. A sample is considered nontoxic at T ≤ 19%, toxic at 19 < T ≤
50%, and strongly toxic at T > 50%.
2.5 Infrared spectra
Infrared spectra were recorded at 400–2200 cm
–1
with a Specord 75IR spectrometer (Karl
Zeiss, Jena, Germany) in KBr pellets (250 mg of KBr + 1.2 mg of a sample). Absorbance
bands were identified according to established methods (Nakanisi, 1965; Sverdlov et al.,
1970; Nakomoto, 1991).
2.6 Gas chromatography/mass spectrometry (GC/MS) analyses
Gas chromatomass spectrometry was performed with a Pegasus 4D chromatomass-
spectrometer (LECO, Russia) under the following conditions: ionization energy 70 eV; 30 m
RTX-5MS capillary silicon column; temperature program 40°C (5 min), 8°C/min, 300°C (10
min); scan range 28–450 Da. Qualitative identification was performed by reference to the
WILEY mass spectrum library, including 270 000 compounds. Perdeuterated naphthalene
was used as an internal reference for quantitative assay.
2.7 The elemental analysis
Analysis of photooxidation products was performed in solutions in distilled water after 13

and 38 h of UV irradiation with air bubbling. Aqueous solutions, tagged as “13" and “38,”
were extracted with dichloromethane 3 times for 10 min each. The extracts were combined,

Herbicides – Environmental Impact Studies and Management Approaches

64
dried with anhydrous sodium sulfate, and concentrated in vacuum to 1 ml. Elemental
analysis was performed after DCPA degradation. Samples were evaporated at 50–80°C, and
C, N, Cl, and H were assayed in the completely dried residue (Klimova, 1975).
The contents of C, H and dry residue were determined by the modified Pregl method (Pregl,
1934) based on the ratio of molecular masses of the elements. A weighed sample of 3-5 mg
was burned in a current of neat oxygen at t = 1040
0
C. The amount of formed CO
2
and H
2
O
was determined by weighing the corresponding absorbers filled with ascarite (NaOH) and
anhydrone (anhydrous Mg perchlorate – Mg(ClO
4
)
2
).
The content of N was determined by the Pregl-Dumas method (Steuermark, 1961). A
weighed sample of 3-5 mg was burned in a current of CO
2
in the presence of CuO and then
reduced by metallic copper. The volume of evolved gaseous nitrogen was determined with
a gas meter.

The content of Cl was determined according to the Schöniger flask method (Schöniger,
1955). The analyzed substance (2-4 mg) was wrapped in filter paper placed in a Pt grid and
hanged up to the cork of the flask. The flask was filled with oxygen. 2N KOH (1 ml), 0.5 мл
H
2
О
2
(0.5 ml), and bidistilled H
2
О (5 ml) were placed on the flask bottom. The end of the
paper (in which the weighed sample is wrapped) was ignited, and the flask was rapidly
corked. Absorption was carried out for 0.5-1 hour. The contents was titrated with 0.01N
Hg(NO
3
)
2
in the presence of diphenylcarbazone to lilac-violet color. The content of Cl was
determined by the titrant volume.
3. Results and discussion
Prior to the experiment, nine lower species were identified in the AS sample under study:
algae, amoebae, sessile infusoria, and flagellates (Table 1). Addition of the pollutant reduced
the number of species to seven. One group, blue-green algae, became predominant,
apparently being the most resistant. Unicellular algae have also been reported as resistant to
the herbicides Diuron and o-Phenanthroline (Laval-Martin et al., 1977).
Treatment with the mutagen caused a change in the species composition of AS and an
increase in the range of the species. This increase was related to the fact that the populations
of some species were too small to be detected before addition of NMU or DCPA. The
presence of a certain pollutant may provide conditions for accelerated growth of those
species that consume the pollutant as a preferable nutrient, thus allowing their
identification. This phenomenon provides grounds for the purification of wastewaters from

chemicals (Golubovskaya, 1978). In our experiment, the biocenosis that formed after NMU
treatment included two phyla of lower plants (bacteria and algae) and five invertebrate
classes (ciliates, sarcods, mastigiophores, nematodes, and rotifers). The number of species
identified in AS sample 3, treated with NMU for 18 h, was greater than in sample 2, treated
for 6 h (20 and 15 species, respectively). The population of rotifers notably increased, and
sulfur bacteria and testaceous amoebae were identified (Table 1). This community
successfully resisted the anthropogenic load. The toxicity of the sample treated for 18 h was
lower than that after 6-hour treatment (Fig. 1). Long-term exposure (several months)
reduced the range of species.

Investigation of Degradation of Pesticide Lontrel in Aqueous Solutions

65
Organism
Control
(without
NMU
treatment)
NMU
treatment,
6 h
NMU
treatment,
18 h
Repeated
treatment,
6 h
Repeated
treatment,
18 h

Ulothrix sp.
 
mass
 
Scenedesmus
obliguus
   

Chlorella vulgaris



Flagellata sp.

mass
Oicomonas socialis

 
Bodo globosus

   
Zooglea ramigera
occasional
Euglena viridis


Filamentous
bacteria
occasional
Bacillus


Beggiatoa minima




Jromia neglecta sp.
 

Arcella vulgaris



Centropixis acullata




Pamphagus hyalinus
 


Euglypha laevis



Amoeba sp.
 
Aspidisca costata




Aspidisca lynceus

 

Lacrimaria sp.




Litonotus anser





Chilodonella uncinata



Vorticella alba



Thuricola similes


Telotpox



Rotaria rotatoria

mass



Colurella s
p
.




Rotaria neptunia


Lelane Monostyla



Brachionus angularis


Cephalodella gibba




Notommata sp.



Cephalodella forticula


Chaetonotus
brevis
p
inosus




Table 1. Hydrobiological analysis of activated sludge.
The presence of DCPA, its copper complex, or industrial waste (samples 1–5) at
concentrations used in the experiment exerted acute and chronic effects on the infusorium
culture. Figure 1 shows that the toxicity of sample 1 remained high (~80%) after two months

Herbicides – Environmental Impact Studies and Management Approaches

66
of monitoring. The toxicity of sample 2, treated with NMU for 6 h, remained high (~90%) for
36 days. Then, it decreased abruptly, and, after 56 days, the sample was virtually nontoxic.
By the beginning of the third month, its toxicity reached 70%. As mentioned above, the
decrease in toxicity after an 18-h treatment (as compared to the 6-h exposure to NMU) may
be related to a change in the proportions of species in AS as a result of the mutagenic effect.
The history of toxicity of sample 4 suggests that the CuL

2
had no pronounced toxic effect on
the AS for 20 days after the onset of exposure to NMU. This result may be related to a
decrease in the reactivity of DCPA in the copper complex. After 20 days, the complex
appears to have been degraded and the toxicity of the sample abruptly increased. The
history of toxicity of sample 5, containing industrial waste, was similar to that of sample 4.
This finding suggests the presence of various complexes between DCPA and metals, whose
decay after 20 days increases the toxicity of the sample dramatically.
0 122436486072
0
20
40
60
80
100
5
4
3
2
1
%
time (days)

Fig. 1. Changes in the toxicity (with B. harveyi) of AS samples containing (1, 2, 3) DCPA, (4)
the CuL
2
complex, and (5) industrial waste water: (1) control sample with intact AS, (2) AS
treated with NMU for 6 h, and (3) AS treated with NMU for 18 h. Starting concentrations:
DCPA, 1.22·10
–3

M; industrial wastewater, ≈ 1·10
–3
M; the CuL
2
complex, 0.4·10
–3
M.
Temperature 25°C. Arrows indicate repeated NMU treatment.
As seen in Fig. 2, DCPA concentrations in samples 4 and 5 remained constant throughout
the experiment. Most likely, it is only the proportions of various CuL
2
that changed, thereby
altering the toxicity of these samples (Fig. 1). In samples 2–4, AS was treated with NMU to
obtain mutations most resistant to the toxic substance under study. After the first 20 days
(Fig. 2), DCPA concentrations changed neither in the control sample nor in samples exposed
to the mutagen. After 18–20 days, DCPA concentrations in samples 1–3 began to decrease.
The decrease in DCPA concentrations in samples 2 and 3 occurred faster than in the control
sample. A steady state was established approximately 40 days after the beginning of the

Investigation of Degradation of Pesticide Lontrel in Aqueous Solutions

67
experiment. Another treatment with NMU was undertaken to further increase the oxidative
potential of AS; however, this attempt did not cause a significant degradation of the
substances under study. After 66 days, the content of DCPA decreased by 25%, while that in
the sample not treated with NMU decreased by 20%.
0 12 24 36 48 60 380 381
0,0
0,2
0,4

0,6
0,8
1,0
1,2
1,4
4
2,3
1
5
10
3
, М
time (days)

Fig. 2. Kinetic curves of the degradation of (1, 2, 3) DCPA, (4) the CuL
2
complex, and (5)
industrial wastewater. Designations follow Fig. 1.
After one year of treatment of the pollutants with a solution of AS treated with the mutagen,
DCPA concentration had decreased by 45% in total, i.e., by less than half; in the control
(nonmutagenized) AS sample, the DCPA concentration had decreased by 30% in total.
During the same one-year term, the concentration of toxic substances in sample 5,
containing wastewater of the herbicide mixture, had decreased only by 7% in total.
Thus, our data indicate that DCPA the main active principle of herbicide Lontrel belongs to
bioresistant organochlorine herbicides. Wastewater treatment in industrial treatment plants
lasts only for 8 ÷ 11 h. During this treatment, Lontrel can form compounds with copper or
other industrial pollutants. It is natural to assume that both Lontrel and its metal derivatives
remain intact in treatment plants. Hence, the toxic substances can penetrate surface or
subsurface water bodies and accumulate there, posing a hazard for microflora, plants, and
fish. Because of Lontrel bioresistance, its application in agriculture for controlling weeds can

exert selection pressure and give rise to mutant weeds, resistant to the herbicide. This effect
has been shown for Simazine and Atrazine, which are resistant to biodegradation in soil
(Fedtke, 1985). Bioresistance was the reason for prohibition of DDT, although it was
nontoxic at recommended concentrations (Burgelya & Myrlyan, 1985).

Herbicides – Environmental Impact Studies and Management Approaches

68
0 4 8 12 16 20 24 28
0
20
40
60
80
100
concentration (%)
time (h)
5
4
6
3
2
1

Fig. 3. Changes in DCPA concentrations (as percentages of the starting concentration), as
determined by the time of UV irradiation: (1) without bubbling; (2) argon, DRB-8 lamp; (3)
air, DRB-8 lamp; (4) ozone, DRB-8 lamp; (5) air, DRSh-1000 lamp; and (6) oxygen, hydrogen
peroxide, BRA-15 lamp.
The high resistance of Lontrel to biodegradation under natural sunlight prompted us to
perform an experiment on photochemical oxidation of the herbicide with the use of

hardUV radiation, which is the only part of the solar UV spectrum (wavelength range,
400–360 nm) that reaches the surface of Earth, shorter waves being trapped by the
atmosphere. Photochemical activity can be manifested by radiation within the absorption
spectrum of a compound. In our case, for DCPA, the active wavelength is  = 283 nm.
Therefore, we used mercury lamps with emission lines within 254–579 nm. The energy of
a quantum with a wavelength of 283 nm exceeds the energy of the C=N bond (~260
kJ/mol) by a factor of about 1.5, an amount of energy that may be sufficient for cleavage
of the very stable pyridine ring, which is not oxidizable by AS microorganisms. The
efficiency of UV irradiation in inducing redox degradation of various contaminants has
been shown previously (Skurlatov & Shtamm, 1997b; Kruithof et al., 1992). It has been
reported that Atrazine is degraded by UV irradiation alone within 20 min and by ozone
alone within almost 3 h (Prado & Esplugas, 1999). Application of UV radiation, including
its combination with the use of hydrogen peroxide, to the degradation of phenols has
been reported (Legrini et al., 1993; Guittonneau et al., 1988; Sundstrom et al., 1989;
Castrantas & Gibilisco, 1990).